Biogeochemical and kinetics characterization of sulfate reducing microbial communities enriched from mine drainages

of sulfate reducing microbial communities

enriched from mine drainages

By

Karabelo Macmillan Moloantoa February 2015

University of the Free State

Universiteit van die Vrystaat

ii

of sulfate reducing microbial communities

enriched from mine drainages

By

Karabelo Macmillan Moloantoa

BSc. Hons. (UFS)

Submitted in fulfilment of the requirements for the degree

MAGISTER SCIENTIAE

In the

Department of Microbial, Biochemical and Food Biotechnology Faculty of Natural and Agricultural Sciences

University of the Free State Bloemfontein

South Africa

February 2015

Supervisor: Prof. E. van Heerden

Co-Supervisor: Dr. J. C. Castillo

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I hereby dedicate this dissertation to my mother

Judith Letladi Moloantoa, Grandmother Eva Tlou Ramphele,

my two uncles (Christopher Ramphele and Lebogang Ramphele), my

three aunts (Mathea Maponya, Phuti Machaba and Katlego Thobejane)

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“With men, this is impossible, but with God all things are possible”

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I would like to express my gratitude and appreciation to the following:

 God: My creator and father, for giving me life, strength and wisdom to accomplish all tasks presented to me during my studies. (Phillipians 4:19 – 20).

 Prof. E. van Heerden: My supervisor, for being the best study leader ever. Thank you for believing in me and never giving up on me even when I was at my worst lows. Thank you for the opportunity to explore my potential in achieving greater things in my life. Your amazing knowledge, assistance, guidance, love and coaching has impacted my life not only academically but personally and intellectually as a scientist.

 Dr. J. C. Castillo: My co-supervisor, for your hard work and guidance throughout my bench work and dissertation write up. Thank your for sacrificing most of your personal time including seval evenings we spent in the laboratory. Thank you for being patient with me and always coming to my rescue when I needed scientific help. I highly value and appreciate your support and guidance till the very last day of the preparation of this script.

 Prof. M. Smith, L. Steyn and T. Mpeyakhe (Department of Biotechnology,

UFS, Bloemfontein): for assistance with the Bioreactors.

 Prof. P. van Wyk, Dr. C. Swart-Pistor and H. Grobler (Centre for

Microscopy, UFS, Bloemfontein): for assistance with SEM and TEM

analysis.

 D. Welman-Purchase (Department of Geology, UFS, Bloemfontein): for assistance with XRD analysis.

vi  Dr. P. Williams, M. Maleke, C. Mulandu, O.O. Kuloyo, A.O. Ojo: My colleagues, for your assistance when I needed extra pair of hands, eyes and scientific knowledge you imparted during my studies and for proofreading my dissertation in you inconvenient times.

 L.J. Moloantoa: My mother, for allowing me to further my studies and supporting me throughout my entire life. I wouldn’t have done it without you. I love you mom.

 Pastor. At Boshof and the CRC pastoral and leadership team (CRC

Bloemfontein): My spiritual leaders, for the spiritual and moral design in my

life. Thank you for your prayers and spiritual guidance and teachings than moulded the person I am today.

 Lesego Sebatana: For the support, genuine love and for always being there for me when I needed someone to confide in. You are the best.

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I hereby declare that this dissertation is submitted by me for the Magister Scientiae degree at the University of the Free State. This work is solely my own and has not been previously submitted by me at any other University or Faculty, and the other sources of information used have been acknowledged. I further grant copyright of this dissertation in favour of the University of the Free State.

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LIST OF ABBREVIATIONS ... xiv

LIST OF FIGURES ... xvii

LIST OF TABLES ... xxiii

CHAPTER 1

1. LITERATURE REVIEW ... 2

1.1. Introduction ... 2

1.2. Acid Mine Drainage ... 2

1.2.1. Background of Acid Mine Drainage ... 2

1.2.2. Acid Mine Drainage formation ... 3

1.2.3. Properties of Acid Mine Drainage ... 5

1.3. Acid Mine Drainage treatment options ... 7

1.3.1. Overall Acid Mine Drainage treatment options ... 7

1.3.2. Chemical Acid Mine Drainage treatment ... 8

1.3.3. Biological Acid Mine Drainage treatment ... 9

1.4. Role of Sulfate-reducing bacteria in Acid Mine Drainage ... 10

1.4.1. Characteristics and diversity of sulfate-reducing bacteria ... 11

1.4.2. Factors affecting biological sulfate reduction in AMD ... 14

1.4.2.1. Electron donor and carbon source selected for SRB growth ... 14

1.4.2.2. The effect of temperature on SRB... 15

1.4.2.3. The effect of pH on sulfate reduction ... 16

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1.5. Conclusions... 20

1.6. References ... 21

CHAPTER 2

2. INTRODUCTION TO THE PRESENT STUDY ... 28

2.1. Introduction ... 28

2.1.1. Study aims and objectives ... 29

2.2. References ... 31

CHAPTER 3

3. THE STUDY AND EVALUATION OF MICROBIAL COMMUNITIES FROM ACID MINE DRAINAGES AND NON ACID MINE DRAINAGES FOR THE SULFATE REDUCTION POTENTIAL ... 33

3.1. Abstract ... 33

3.2. Introduction ... 34

3.2.1. Aims and objectives of the chapter ... 36

3.3. Materials and methods ... 37

3.3.1. Localization and environmental settings ... 37

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3.3.3.1. Physicochemical parameters ... 39

3.3.3.2. Sulfate concentrations ... 40

3.3.3.3. Sulfide concentrations ... 40

3.3.3.4. Other chemical analysis ... 42

3.3.4. Microbial characterization ... 42

3.3.4.1. DAPI staining ... 42

3.3.4.2. Live/Dead staining... 44

3.3.4.3. Gram staining ... 44

3.3.5. Molecular characterization ... 45

3.3.5.1. Genomic DNA extraction ... 45

3.3.5.2. Amplification of the partial 16S rRNA gene fragment ... 46

3.3.5.3. Denaturation Gradient Gel Electrophoresis (DGGE) ... 47

3.3.6. Microbial enrichments ... 51

3.3.6.1. Anaerobic media preparations ... 52

3.3.6.2. Media inoculation and incubation ... 53

3.3.6.3. Secondary inoculation into fresh media ... 54

3.3.6.4. Tertiary inoculation into the fresh media ... 54

3.3.6.5. Microbial characterization of enriched cultures ... 55

3.3.6.6. Molecular analysis of the enriched tertiary cultures... 55

3.3.6.7. Amplification of the dsrAB gene fragments from gDNA... 55

3.3.6.8. Cryopreservation of the enriched cultures ... 57

3.3.6.9. Scanning Electron Microscopy ... 57

3.4. Results and discussions ... 58

3.4.1. Samples description and analysis ... 58

3.4.2. Microbial characterization results ... 61

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3.4.2.3. Morphological characterization of bacteria by gram staining ... 62

3.4.3. Molecular characterization results ... 63

3.4.3.1. Genomic DNA extraction ... 63

3.4.3.2. Amplification of the partial 16S rRNA gene fragments ... 65

3.4.3.3. Denaturation Gradient Gel Electrophoresis (DGGE) analysis ... 65

3.4.4. Microbial enrichments results ... 70

3.4.4.1. Results analysis of the anaerobic enrichments ... 70

3.4.4.2. Microbial characterization of tertiary enrichment cultures ... 71

3.4.5. Molecular analysis of the enriched tertiary cultures ... 72

3.4.5.1. Genomic DNA extraction of the tertiary enrichment cultures ... 72

3.4.5.2. Amplification results of the partial 16S rRNA fragments from the enrichments gDNA ... 73

3.4.5.3. Denaturation Gradient Gel Electrophoresis (DGGE) analysis of the enriched cultures ... 74

3.4.5.4. Amplification results of the dsr gene fragments from the extracted gDNA 77 3.4.5.5. Scanning Electron Microscopy analysis of the enriched cultures ... 79

3.5. Conclusions... 82

3.6. References ... 84

CHAPTER 4

4. BEHAVIOURAL STUDIES AND EVALUATION OF ANAEROBIC SULFATE REDUCING COMMUNITIES FOR SULFATE REDUCTION AND METAL PRECIPITATION PROCESSES IN VARIED CONDITIONS ... 91

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4.2.1. Aims and objectives of the chapter ... 94

4.3. Materials and methods ... 95

4.3.1. Preparation and maintenance of cultures for the microbial-metal interaction experiments ... 95

4.3.2. Batch experiments ... 96

4.3.2.1. Microbial sulfate reduction of inclining concentrations ... 96

4.3.2.2. Effects of pH and temperature on sulfate reduction ... 96

4.3.2.2.1. Optical density measurements ... 97

4.3.2.3. Effects of high metal concentrations on sulfate reduction ... 97

4.3.3. Bioreactor studies ... 98

4.3.3.1. Carbon source selection ... 98

4.3.3.1.1. The Sixfors bioreactor setup ... 98

4.3.3.1.2. Bioreactors inoculation, start up, operation and termination ... 100

4.3.3.2. Metal-microbe interactions studies in the bioreactors ... 101

4.3.3.2.1. Sampling and analysis ... 101

4.3.3.2.2. Carbon source analysis ... 102

4.3.3.3. Molecular analysis of metals tolerant microorganisms ... 103

4.3.3.4. Morphological and mineralogical characterization of precipitates 103 4.3.3.4.1. Samples preparation ... 103

4.3.3.4.2. Scanning Electron Microscopy coupled to Energy Dispersed X-ray Spectroscopy (SEM-EDS) ... 103

4.3.3.4.3. Transmission Electron Microscopy coupled to Energy Dispersed X-ray Spectroscopy (TEM-EDS) ... 104

4.3.3.4.4. X-Ray Diffractometer analysis (XRD) ... 104

4.4. Results and discussions ... 105

4.4.1. Culture preparations for the metal-microbe interaction ... 105

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4.4.2.2. Effects of low pH and temperature on sulfate-reduction ... 107

4.4.2.3. Effects of high metal concentrations on sulfate reduction ... 111

4.4.3. Bioreactor studies ... 114

4.4.3.1. Carbon source selection ... 114

4.4.3.2. Metal-microbe interactions ... 117

4.4.3.3. Molecular analysis of metals tolerant microorganisms ... 122

4.4.3.3.1. Genomic DNA extraction and 16S rRNA amplification ... 122

4.4.3.3.2. Denaturation Gradient Gel Electrophoresis (DGGE) ... 123

4.4.3.4. Morphological and mineralogical characterization of precipitates 129 4.4.3.4.1. Scanning Electron Microscopy coupled to Energy Dispersed X-ray Spectroscopy (SEM-EDS) ... 130

4.4.3.4.2. Transmission Electron Microscopy coupled to Energy Dispersed X-ray spectroscopy (TEM-EDS) ... 134

4.4.3.4.3. X-Ray Diffraction analysis ... 136

4.5. Conclusions... 137 4.6. References ... 139

CHAPTER 5

xiv ºC Degrees Celsius < less than > greater than % Percentage µL microlitre

AMD acid mine drainage

ARD Acid rock drainage

ASRM anaerobic sulfate reducing medium

BLAST Basic Local Alignment Search Tool

bp base pair

DAPI 4’,6-diamidino-2-phenylindole

DGGE Deanaturation Gradient Gel Electrophoresis

dH2O distilled water

DNA Deoxyribonucleic Acid

DO Dissolved oxygen

DSR Dissimilatory sulphite reductase enzyme

dsrAB Dissimilatory sulfite reductase gene

EC Electrical conductivity

EDS Energy Dispersive X-ray Spectroscopy

EDTA Ethylenediaminetetraacetic acid.

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et al. et alii / and others

EtBr Ethidium Bromide

FeT Total iron

g gram

gDNA Genomic DNA

ha hectare

HPLC High Performance Liquid Chromatography

ICP Inductively Coupled Plasma

IDT Integrated DNA Technologies

IGS Institute of Ground water Studies

km kilometre

kV kilovolts

L litre

M Molar

mg/L milligram per litre

mL millilitre

mM millimolar

mS/m Millisimens per metre

mV millivolts

NCBI National Centre for Biotechnology Information

ng/µL nanogram per microlitre

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OD Optical Density

ORP Oxidation Reduction Potential

PCR Polymerase Chain Reaction

pH Measure of the Acidity or Basicity of a solution

PSGM Postgate medium

RNA Ribosamal nucleic acid

SABS South African Bureau of Standards

SEM Scanning Electron Microscopy

SOB Sulfide-oxidizing bacteria

sp. specie

spp. species

SRB sulfate-reducing bacteria

t ton

TAE Tris Acetate EDTA

TDS Total Dissolved Solids

TEM Transmission Electron Microscopy

UF Urea Formamide

v/v volume per volume

w/v weight per volume

w/w weight per weight

x g Acceleration due to Gravity

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Figure 1.1: Biological (biotic) and chemical (abiotic) strategies for acid mine drainage remediation. (Taken from Johnson and Hallberg, 2005).

Figure 1.2: Biological sulfur transformations in natural and artificial environments. (Taken from Sánchez-Adrea et al., 2014).

Figure 1.3: Sulfide speciation as a function of pH at 25ºC. (Taken from Kaksonen and Puhukka, 2007).

Figure 3.1: AMD and NMD chemical formation process and the microbial diversities dominant in them. (Generic figure).

Figure 3.2: South African map indicating selected study sites. Location 1: Site-Ex, Location 2: Site-Ka and Location 3: Site-Po (Taken from McCarthy, 2011).

Figure 3.3: Sampling sites selected for the study. A: Site-Ex, B: Site-Ka and C: Site-Po.

Figure 3.4: Standard curve relating sulfide concentrations to absorbance (670 nm). Figure 3.5: Anaerobic serum vials containing enrichment media. Left: PSGM,

Middle: two vials containing yeast extract and Right: ASRM.

Figure 3.6: DAPI staining images. A: Sludge sample, B: Site-Ex, C: Site-Ka and D: Site-Po. Scale bar = 1 µm.

Figure 3.7: Live/Dead staining images from the three water samples. A: Site-Ex, B: Site-Ka and C: Site-Po.

Figure 3.8: Gram stained bacterial cells from the three water samples. A: Site-Ex, B: Site-Ka and C: Site-Po.

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samples.

A. Lane M: DNA standard Marker (O’GeneRulerTM Ladder Mix), Lane 1: Site-Ex, Lane 2: Site-Ka and Lane 3: Site-Po.

B. Lane M: DNA standard Marker (O’GeneRulerTM Ladder Mix), Lane 4: Sludge 1 and Lane 5: Sludge 2.

Figure 3.10: Amplification results of the partial 16S rRNA fragments from the extracted gDNA on a 1% (w/v) agarose gel. Lane M: DNA standard Marker (O’GeneRulerTM

Ladder Mix), Lane 1: Sludge 1, Lane 2: Sludge 2, Lane 3: Site-Ex, Lane 4: Site-Ka, Lane 5: Site-Po and Lane 6: Positive control.

Figure 3.11: DGGE profile of the sludge and AMD water samples. Lane 1: Site-Ex, Lane 2: Site-Ka, Lane 3: Site-Po and Lane 4: Sludge sample.

Figure 3.12: Tertiary anaerobic enrichment cultures in both ASRM and PSGM. A1: Sludge in ASRM, A2: Site-Ex in ASRM, A3: Site-Ka in ASRM and A4: Po in ASRM. B1: Sludge in PSGM, B2: Ex in PSGM, B3: Site-Ka in PSGM and B4: Site-Po in PSGM.

Figure 3.13: Microscopic analysis results of tertiary enrichment cultures. A-D: Gram staining (PSGM). A: Sludge in PSGM, B: Site-Ex in PSGM, C: Site-Ka in PSGM and D: Site-Po in PSGM.

E-H: Gram staining (ASRM). E: Sludge in ASRM, F: Site-Ex in ASRM, G: Site-Ka in ASRM and H: Site-Po in ASRM.

I-L: DAPI staining. I: Sludge in PSGM, J: Site-Ex in PSGM, K: Site-Ka in PSGM and L: Site-Po in PSGM.

M-P: Live/Dead staining. M: Sludge in PSGM, N: Site-Ex in PSGM, O: Site-Ka in PSGM and P: Site-Po in PSGM. Scale bar = 1 µm.

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DNA standard Marker (O’GeneRulerTM

Ladder Mix), Lane 1: Sludge in PSGM, Lane 2: Sludge in ASRM, Lane 3: Site-Ex in PSGM, Lane 4: Site-Ex in ASRM, Lane 5: Site-Ka in PSGM, Lane 6: Site-Ka in ASRM, Lane 7: Site-Po in PSGM and Lane 8: Site-Po in ASRM.

Figure 3.15: DGGE profile of tertiary cultures enriched in ASRM and PSGM. Lane 1: Sludge in PSGM, Lane 2: Sludge in ASRM, Lane 3: Site-Ex in PSGM, Lane 4: Site-Ex in ASRM, Lane 5: Site-Ka in PSGM, Lane 6: Site-Ka in ASRM, Lane 7: Site-Po in PSGM and Lane 8: Site-Po in ASRM.

Figure 3.16: Amplification results of the dsrAB gene fragments on 1% agarose gels. A: PCR products of the dsrAB gene fragments. Lane M: DNA standard Marker (O’GeneRulerTM

Ladder Mix), Lane 1: Site-Ex in PSGM, Lane 2: Site-Ka in PSGM, Lane 3: Slidge in PSGM, Lane 4: Sludge in ASRM, Lane 5: Site-Po in ASRM.

B: Gradient PCR amplification of the dsrAB gene fragments from sludge samples. Lane M: DNA standard Marker (O’GeneRulerTM

Ladder Mix), Group 1 Lanes: Sludge in PSGM and Group 2 Lanes: Sludge in ASRM.

Figure 3.17: Scanning Electron Microscope image showing biofilm and precipitates formed in the tertiary enrichment cultures. A: Biofilm enriched in PSGM, B and C: Higher magnification of subsections of A, D: Biofilm enriched in ASRM, E and F: Back Scattering images of precipitates in culture enriched in PSGM, G: Back Scattering images of precipitates in culture enriched in ASRM, H: Reference image of Pyrite precipitates (Taken from Larrasoaña et al., 2014), I: Reference image of the

Desulfovibrio sp. (Taken from Warthmann et al., 2005).

Figure 4.1: Preparation flow diagram and selected parameters for evaluating sulfate reducing activity of the enriched SRB.

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Figure 4.3: HPLC standard curve relating glycerol peak area to the amount of sample injected into the HPLC column.

Figure 4.4: Live/Dead staining images (A to C): A: Co-cultured consortia in PSGM, B: Co-cultured consortia in ASRM and C: Secondary co-cultured consortia from cultures depicted in A and B. D: Consortia maintained in PSGM for downstream experiments.

Figure 4.5: Parameter profiles monitored in four batch experiments to evaluate the effects of pH and temperature on SRB. A: Microbial growth curves, B: ORP profiles, C: pH profiles, D: sulfate reduction profiles, E: sulfide formation profiles and F: sulfide formation profiles of cultures grown at a pH of 3.5 (25ºC) and two cultures both grown at temperature of 10ºC in pH of 6.5 and 3.5.

Figure 4.6: Vials containing cultures used in batch experiments of the metal effects on sulfate reduction and profiles of monitored parameters. A: Vials containing cultures at day 12. A1: Culture with no metals, A2: Culture with 200 mg/L of Fe2+, A3: Culture with 100 mg/L of Fe2+ and 100 mg/L of Zn2+, A4: Culture with 100 mg/L of Zn2+ and A5: Culture with 200 mg/L of Zn2+. B: pH profiles, C: sulfate reduction profiles and D: sulfide formation profiles.

Figure 4.7: Profiles showing evolutions of parameters yielded by two cultures with different carbon sources: glycerol and sodium lactate. A: sulfate reduction profiles, B: sulfide formation profiles, C: ORP profiles and D: DO profiles.

Figure 4.8: Profiles showing evolutions of parameters yielded by four cultures with different metal concentrations and pH control effects on bacterial growth. A: sulfate reduction profiles, B: sulfide generation profiles, C: ORP profiles, D: pH profiles, E: Zn2+ reduction profile, F: Fe2+ reduction

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Figure 4.9: Amplificons of the 16S rRNA fragments from the extracted gDNA of bioreactor samples on a 1% (w/v) agarose gel stained with EtBr. Group 1: R1 samples, Group 2: R2 samples, Group 3: R3 samples and Group 4: R4 samples. Lane M: DNA standard Marker (O’GeneRulerTM

Ladder Mix), Lane 1: day 5, Lane 2: day 10, Lane 3: day 15, Lane 4: day 20 and Lane I: Inoculum used for the bioreactors.

Figure 4.10: DGGE profiles of consortia incubated in reactors with metals (R1: Inoculum, 200 mg/L Zn2+ and pH control and R2: Inoculum, 200 mg/L Fe2+ and pH control).

Group 1 (Fe2+): Lane 1: R1 – day 5, Lane 2: R1 – day 10, Lane 3: R1 – day 15, Lane 4: R1 – day 20.

Group 2 (Zn2+): Lane 5: R2 – day 5, Lane 6: R2 – day 10, Lane 7: R2 – day 15, Lane 8: R2 – day 20.

Figure 4.11: DGGE profiles of consortia incubated in reactors with (R3: Inoculum, no metals and no pH control and R4: Inoculum, no metals and pH control).

Group 1 (-pH): Lane 1: R3 – day 5, Lane 2: R3 – day 10, Lane 3: R3 – day 15, Lane 4: R3 – day 20.

Group 2 (+pH): Lane 5: R4 – day 5, Lane 6: R4 – day 10, Lane 7: R4 – day 15, Lane 8: R4 – day 20.

Figure 4.12: Centrifuged samples before lyophilisation. A: R1 (Inoculum, 200 mg/L Zn2+ and pH control), B: R2, (Inoculum, 200 mg/L Fe2+ and pH control), C: R3 (Inoculum, no metals and no pH control) and D: R4 (Inoculum, no metals and pH control).

Figure 4.13: SEM micrographs and EDS spectrum of sample harvested from a bioreactor with 200 mg/L of Zn2+ (R1). A and B: SEM micrographs showing bacterial cells, C: magnified red square in B and D: SEM-EDS analysis spectrum of the circled particle.

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magnified red square in B and D: SEM-EDS analysis spectrum of the circled particle in C.

Figure 4.15: SEM micrographs of samples harvested from bioreactors with no metals. A to C: SEM micrographs showing bacterial biofilm from the bioreactors (R3 and R4).

Figure 4.16: TEM micrographs and EDS spectrum of Zn2+ containing sample from a bioreactor with 200 mg/L of Zn2+ (R2). A and B: TEM micrographs, C: TEM-EDS analysis spectrum of the encircled area in red.

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Table 1.1: Classification of representative sulfate-reducing bacteria (Taken from Castro et al., 2000).

Table 3.1: PCR primer set and sequences for 16S rRNA amplification. Table 3.2: PCR programme for 16S rRNA amplification.

Table 3.3: Sequencing PCR programme.

Table 3.4: Media compositions of ASRM and PSGM. Table 3.5: Designated names of the enrichment samples. Table 3.6: Sequences of dsrAB gene fragments primer set

Table 3.7: PCR programme for the amplification of dsrAB gene fragments. Table 3.8: Physicochemical characteristics of mine drainages.

Table 3.9: The hydro-geochemical data of the three water samples. Table 3.10: Microbial cell count estimations by DAPI staining technique.

Table 3.11: Concentrations and purity rations of the extracted gDNA from raw samples.

Table 3.12: Sequencing results obtained from BLAST analysis for AMD water samples and the sludge sample

Table 3.13: Concentrations of extracted gDNA from the tertiary cultures. Table 3.14: Sequencing results obtained from BLAST algorithm for tertiary

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Table 4.2: Sulfate reduction efficacy of SRB consortium.

Table 4.3: Sequencing results obtained from the BLAST analysis for samples obtained from bioreactors with metals (R1 and R2).

Table 4.4: Sequencing results obtained from the BLAST analysis for samples obtained from bioreactors with and without pH control (R3 and R4). Table 4.5: SEM-EDS results showing elemental composition of the precipitates in

the Zn2+ sample (R1).

Table 4.6: SEM-EDS results showing elemental composition of the precipitates in the Fe2+ sample (R2).

Table 4.7: TEM-EDS results showing elemental composition of the precipitates in the Zn2+ containing sample analysed on spectrum.

Chapter 1 1

CHAPTER 1

Chapter 1 2

1. LITERATURE REVIEW

1.1. Introduction

Water is an essential element of life and all organisms are dependent on it to exist. It is the most abundant natural resource on earth, however, only 0.016% is liquid fresh water readily available for human consumption and domestic use (Derakhshani and Alipour, 2010; Meyer and Casey, 2004). Pollution of water is a global challenge and most human activities and industries contribute largely to the pollution (Mačingová and Luptáková, 2010). The most dominant water pollutant faced globally is Acid Mine Drainage (AMD) from the active and or abandoned mines which are continuously degrading the quality of groundwater, streams, rivers and complete river basins in mining practising countries (Nieto et al., 2007).

AMD contaminated water is toxic due to the dissolved heavy metals making it domestically un-usable even for irrigation or feedstock consumption. Approximately one ton of water is used for one ton of rock to be crushed and processed to extract valuable minerals making water demands high for large scale mineral mines (White, 1985). The wastewater from the mines affects the groundwater systems and is released on to the surface which tends to flow into streams and rivers (McCarthy, 2011). If the AMD contaminated water persists on infiltrating the fresh streams without any treatment, water demands will continue rising leading to the economical imbalance as more money will be required in future to treat AMD contaminated water in the rivers.

1.2. Acid Mine Drainage

1.2.1. Background of Acid Mine Drainage

Acid mine drainage (AMD) or acid rock drainage (ARD) is a consequence of most industrial activities, principally mining. It is mostly characterized by a low pH, high concentrations of sulfate and heavy metals (Moosa et al., 2005). AMD mostly occurs as runoff or seepage from waste mineral stockpiles or tailings from the mines. The source of AMD is the host rock which contains metal sulfides like pyrite (marcasite,

Chapter 1 3

FeS2), sphalerite (ZnS), galena (PbS), chalcocite (Cu2S), millerite (NiS), cinnabar

(HgS), arsenopyrite (FeAsS) and chalcopyrite (CuFeS2) that are often mined for

extraction of precious metals such as gold (Au), silver (Ag), lead (Pb), iron (Fe), zinc (Zn) and copper (Cu) to be used commercially (Baker and Banfield, 2003; Baker et

al., 2003; Kalin et al., 2006). During mining or any other anthropogenic activity that

involve the excavation of rocks, the host rocks exposes metal sulfides which react with air and water forming AMD (Kalin et al., 200). The fragmented rocks, typically still containing metal sulfides are disposed around the mining area or completed construction sites which are referred to as tailings (Doepker and Drake, 1991). In AMD polluted streams, the water tends to have a pH less than 3.5 and dissolved metals which are harmful to most of the aquatic life (Henry et al., 1999). Not only active mines serve as major sources of AMD but abandoned surface and underground mines also produce AMD that can pollute hundreds kilometres of streams in a decade after mining has stopped (Hedin et al., 2005; Hoffert, 1947). AMD can also form naturally through the weathering process where the host rock is exposed. During rainy seasons, AMD form at a slow pace with low toxicity that might continue for years without recognition as it gets diluted (Gray, 1997).

1.2.2. Acid Mine Drainage formation

AMD formation is a well-studied and understood process (McCarthy, 2011) that involves a series of chemical and biochemical reactions that takes place when pyrite (FeS2, fool’s gold or iron di-sulfide) and other metal sulfide minerals are exposed to

water and oxygen (in the air), normally during mining (Nieto et al., 2007). When metal sulfides are oxidised, metals leach into the drainage resulting in high dissolved metal concentrations which makes the drainage toxic. As the sulfur from the metal sulfide minerals gets oxidised, high sulfate concentrations are formed which react with released hydrogen to form sulphuric acid that makes the drainage acidic and further promoting dissolution of heavy metals (Akcil and Koldas, 2006; Moosa et al., 2005).

Oxidation of pyrite or any other metal sulfide by atmospheric oxygen and water takes place according to Equation 1.1, a so-called “initiator reaction” which releases

Chapter 1 4

ferrous iron, sulfate and hydrogen ions in the drainage (Kalin et al., 2006; Küsel, 2003; Nieto et al., 2007). The reaction is slow due to the insolubility nature of pyrite and most metal sulfides associated with the host rock.

Equation 1.1: 2𝐹𝑒𝑆2+ 7𝑂2+ 2𝐻2𝑂 → 2𝐹𝑒2++ 4𝑆𝑂42−+ 4𝐻+

Further oxidation of ferrous iron to ferric iron takes place in the environment according to Equation 1.2 but the oxidation process is not always spontaneous as the conversion of iron species in AMD partially depends on the pH of the environment. However, the presence of various iron oxidizing bacteria in the AMD catalyse the oxidation process of ferric iron (Akcil and Koldas, 2006; Marini et al., 2003). At pH conditions below 4 which are typical for most AMD, ferrous iron is relatively stable especially in the presence of oxygen (Kalin et al., 2006). The presence of acidophilic bacteria such as Acidithiobacillus ferrooxidans and

Leptospirillum ferrooxidans, catalyses and enhances the oxidation process of ferrous

iron to ferric iron by 104 to 106 times compared to the reaction rate in their absence (Doepker and Drake, 1991; Taylor et al., 1984). In contrast, at pH values above 4, ferric iron is biologically oxidized by iron oxidizing bacteria such as Gallionella

ferruginea which are present in AMD with low acidic conditions (Johnson and

Hallberg, 2005).

Equation 1.2: 𝐹𝑒2++ 𝑂2+ 4𝐻+ → 𝐹𝑒3+ + 2𝐻2𝑂

In pH conditions above 4, exposure of ferric iron to fresh water results in the precipitation of ferric hydroxide (Fe(OH)3) so-called “Yellow Boy” while releasing

protons in the drainage according to Equation 1.3. Yellow Boy is visually detected by an ochre deposit at the bottom of the streams (Atkins and Pooley, 1982; Kakooei et

al., 2012; Kalin et al., 2006).

Equation 1.3: 𝐹𝑒3++ 2𝐻2𝑂 → 𝐹𝑒(𝑂𝐻)3+ 3𝐻+

In highly acidic environments (pH < 4), ferric iron is highly soluble and serves as an oxidizing agent with greater affinity for pyrite (Marini et al., 2003). In these conditions, ferric iron reacts spontaneously with pyrite releasing more ferrous iron, sulfate and protons in the drainage according to Equation 1.4. The oxidation process of pyrite by ferric iron is known as the “ferric shunt” which results in more acid production in

Chapter 1 5

AMD. The net chemically and biologically mediated AMD formation reaction outlined in Equation 1.5 summarises the series of equations that takes place resulting in a highly acidic drainage with high sulfate concentrations (Johnson and Hallberg, 2005; Marini et al., 2003). The oxidation process of metal sulfide minerals by ferric ir

.on can be induced directly by A. ferrooxidans biofilm that form extracellular polymeric substances (EPS) which allow attachment of cells to the metal sulfide to allow ferric iron attack of the mineral sulfide to catalyse the oxidation process (Kakooei et al., 2012; Küsel, 2003; Nieto et al., 2007).

Equation 1.4: 𝐹𝑒𝑆2 + 14𝐹𝑒3+ + 8𝐻2𝑂 → 15𝐹𝑒2+ + 2𝑆𝑂42− + 16𝐻+

Equation 1.5: 4𝐹𝑒𝑆₂ + 15𝑂₂ + 14𝐻₂𝑂 → 4𝐹𝑒(𝑂𝐻)₃ + 8𝐻₂𝑆𝑂₄

The accelerated rate of pyrite oxidation by bacterial activity leads to the rapid decrease of pH in the drainage which in return induces dissolution of metals like Cd2+, Co2+, Cr3+, Cu+, Fe2+, Mn2+, Pb2+ and Zn2+ (Henry et al., 1999; Nieto et al., 2007). The flow rate of water over the host rock, contact time of water on rock and temperature plays important roles in the rate and toxicity of AMD during its formation (Mačingová and Luptáková, 2010). In South Africa, the biggest AMD sources in the environment are the gold, iron and coal mining due to the methods used during their extraction while diamond, magnesium, chrome and vanadium mines produces less to non-significant volumes of AMD (McCarthy, 2011).

1.2.3. Properties of Acid Mine Drainage

Mine drainages differ in characteristics based on the composition of their water source and post formation contaminants (metals) associated with ore minerals that leach in the water during mining. Typical drainages from mostly gold mines can have very low pH values (below 4) hence called acid mine drainages (AMD). Mine drainages with higher pH (above 4 to 9) are called alkaline mine drainages (AMD) also referred to as non-acid mine drainage (NMD). The drainage contain neutralizing minerals which reverse the acidity of AMD that resulted from pyrite oxidation and this drainage is often observed in coal mine (Akcil and Koldas, 2006).

Chapter 1 6

Stagnant or stored AMD in oxic dams with high iron concentrations have stabilized acidic conditions due to the cycling of iron species (Fe2+/Fe3+) mediated by acidophilic bacteria, which prevents methanogenic or sulfate reducing activities to occur as they require anoxic environment to take place (Kalin et al., 2006; Küsel, 2003). However, biofilm of sulfate-reducing bacteria (SRB) and methanogens colonise at the base of the catchment within the sediments where minimal sulfate reduction takes place. Such environments has served as a source of most isolated SRB from AMD such as Desulfovibrio sp., Desulfomonas sp., Desulfobacter sp.,

Desulfococcus sp., Desulfotomaculum sp. and Desulfobulbus sp. (Rzeczycka and

Blaszczyk, 2005; Wargin, 2007). The diversity found in AMD includes bacterial groups that are divisions of Proteobacteria, Firmicutes and Acidobacteria. Within the

Proteobcteria group which is the widest studied group of microbial SRB, there are:

α-proteobacteria (Acidiphilum sp.), β-α-proteobacteria (Thiomonas sp.), ϒ-α-proteobacteria (Acidithiobacillu sp. and Thiobacillus sp.) and ᵟ-proteobacteria (Desulfovibrio sp.). From the Archaeal lineage, Thermoplasmatales and Sulfolobales were reported to form part of the microbial diversity in AMD and from the Eucarya lineage, Ciliates (Cinetochilium genus) and amoeba (Vahlkampfi sp.) are the most isolated eukaryotes from AMD (Baker and Banfield, 2003; Kalin et al., 2006; Pini et al., 2011; Quaiser et al., 2003; Sharmin et al., 2013).

NMD is common in most coal mining drainages and formed similar to AMD which result in high sulfate but low metals concentrations due to higher pH values (4 – 9 (Akcil and Koldas, 2006; Marini et al., 2003). NMD also occur in flowing streams contaminated with AMD in which the pH of the water rises as the AMD gets diluted by fresh flowing water but the sulfate concentrations remain relatively high because of the continuous oxidation of tailing while ferric hydroxide precipitates to the bottom of the stream with other metals discolouring the water (Akcil and Koldas, 2006). The water mostly appears blue due to the precipitated metals which discolour the bottom of the streams (Marini et al., 2003). In most coal associated rocks, neutralizing chemicals like sodium carbonate or bicarbonate (Na2CO3 or Na2HCO3) and calcium

carbonate (CaCO3) also known as calcite are present and serve as a buffering

agents by raising the pH of AMD to over 6.5 which is a typical pH for NDM (Akcil and Koldas, 2006; Foti et al., 2007). Due to the higher pH conditions, some metals like Cd2+, Ni2+, Fe2+, Zn2+, Cu2+ and Pb2+ precipitate to the bottom of the streams like

Chapter 1 7

Fe(OH) (Rose et al., 1998; Zagury et al., 2006). However, high concentrations of Ca2+, Mg2+, Na+, SO42- and NO3- remain dissolved in the drainage because the high

pH (> 6.5) conditions do not induce their precipitation. Lower diversity of SRB species and other extremophiles have been detected in NMD with dominant species being Acidiphilium sp. and Gallionella sp. (Atkins and Pooley, 1982; Johnson and Hallberg, 2005; Lin et al., 2012).

1.3. Acid Mine Drainage treatment options

In order to treat mine drainage (AMD or NMD) from any mine, first the process of mobilization, transportation and sequestration of relevant chemical components must be understood and characterized to apply relevant treatment option for the drainage (Marini et al., 2003).

1.3.1. Overall Acid Mine Drainage treatment options

Treating AMD from its source is considered the best way to prevent AMD from reaching the fresh flowing water but this is not always possible for most impacted areas (Johnson and Hallberg, 2005). There are two options of remediating AMD which are chemical and biological systems that employ abiotic and biotic strategies respectively (Sierra-Alvarez et al., 2006). Both strategies have been applied widely to raise the pH and to mitigate heavy metals from the drainages without any further environmental pollution especially by the disposal of the by-products of the treatment option. Further classification of the treatment options was developed and the chemical and biological systems were separately classified into both active and passive treatments according to Figure 1.1 (Johnson and Hallberg, 2005). Passive treatment requires less resources to operate and mostly utilizes natural resources which do not need constant monitoring once in operation whereas active treatment requires constant monitoring and continuous input of not necessarily natural resources once in operation to sustain the process (Johnson and Hallberg, 2005; van Eeden et al., 2009).

Chapter 1 8 Figure 1.1: Biological (biotic) and chemical (abiotic) strategies for acid mine drainage

remediation. (Taken from Johnson and Hallberg, 2005).

1.3.2. Chemical Acid Mine Drainage treatment

Chemical treatments involve addition of alkaline material like limestone (calcium oxide, CaO) directly into the AMD to raise the pH. However, the method results in the formation of slurry metals precipitate which are also toxic (Kalin et al., 2006). Passive systems of abiotic treatment option have always been the first preference for most industries with the aim of cutting maintenance cost of the systems. In passive systems, limestone is added in anoxic environment to avoid ferric iron oxidative reactivity to be activated hence increasing the pH of the drainage. Although passive anoxic treatment is effective to certain extend, it is not suitable to treat all kinds of AMD and the volumes of AMD required to be treated daily (Rose et al., 1998; Kalin

et al., 2006).

Active abiotic systems complement most of the remediation activities unachieved by the passive abiotic system. This employs aeration of the system with simultaneous addition of a chemical-neutralizing agent such as: lime, calcium carbonate (CaCO3),

Chapter 1 9

sodium carbonate (NaCO3), sodium hydroxide (NaOH) and magnesium oxide

(MgOH) (Neculita et al., 2007). The neutralizing agent increases the pH of the AMD which will directly induce a rapid oxidation of ferrous iron that in return precipitate with dissolved metals forming a slurry by-product with high concentrations of metals which will require pre-treatment before disposal (Johnson and Hallberg, 2005; Zagury et al., 2006).

1.3.3. Biological Acid Mine Drainage treatment

Biological systems employ biotic entities such as plants, algae, fungi and bacteria to treat AMD normally in bioreactors or wetlands and the process is called “Bioremediation” (Derakhshani and Alipour, 2010; Rose et al., 1998). Passive biological treatments are carried out in both aerobic and anaerobic wetlands where AMD is channelled through a treatment system containing reactive barriers with biological entities like bacteria, algae and plants that assist in precipitating dissolved metals. Aerobic wetlands are normally used for the mildly acidic or net-alkaline water with elevated iron concentrations. The aeration of the wetland is necessary for dissolved iron to oxidize and precipitate with other metals hence remediating the drainage during the flow (Al-Zuhair et al., 2008). The anoxic passive biological systems require construction of reactors which include the addition of lime stone beneath or mixed with organic substrate to be utilized by supplied inoculum of bacteria like SRB to remediate AMD. This option has been widely used to treat highly acidic AMD with elevated sulfate concentrations (Neculita et al., 2007). The system aims at reducing sulfate by SRB while oxidizing the organic matter to carbonates that generate alkalinity in the drainage according to Equation 1.6 (Neculita et al., 2007). Sulfate-reduction process occurs in anoxic environments with low to no oxygen mediated by SRB. The SRB utilize SO42- as an electron acceptor

converting it into hydrogen sulphide (H2S) that react with metals (Me2+) to form metal

sulfide precipitates (MeS↓) according to Equation 1.7 (Luptakova et al., 2012;

Neculita et al., 2007; Zagury et al., 2006). During this process, sulfate concentrations are reduced and metals are mitigated from the drainage through precipitation (Zagury et al., 2006). One of the advantages of using this method is that the

Chapter 1 10

precipitated metals can be retrieved from the by-products and used for commercial purposes.

Equation 1.6: 𝑆𝑂₄2− + 2𝐶𝐻₂𝑂 → 𝐻₂𝑆 + 2𝐻𝐶𝑂₃− Equation 1.7: 𝐻₂𝑆 + 𝑀𝑒2+ → 𝑀𝑒𝑆_↓+ 2𝐻+

Active biological systems rely entirely on bioremediation to reduce sulfate concentrations and increase the pH of mostly acidic AMD. Microorganisms capable of metals immobilization like SRB are employed in anoxic sulfidogenic bioreactors to remediate AMD (Johnson and Hallberg, 2005). The anaerobic bioreactors constructed for this type of treatment act as reversal chambers of AMD formation process that takes place when alkalinity is generated in the environment with subsequent metals and sulfate reduction (Sierra-Alvarez et al., 2006). In the bioreactors with bacterial consortia aimed at remediating AMD, processes such as methanogenesis, fermentation, denitrification, iron reduction, manganese reduction, ammonification and sulfate reduction takes place depending on the chemical composition of the drainage (van Eeden et al., 2009). Organic substrate is required in the bioreactor operations to serve as the carbon source and electron donor needed for bacterial activities that reverse the AMD acidity and toxicity by immobilizing metals. To acquire best remediation results from severe cases of AMD contamination, both passive and active biological treatments are applied respectively but require constant monitoring and maintenance (Sierra-Alvarez et al., 2006; Zagury

et al., 2006).

1.4. Role of Sulfate-reducing bacteria in Acid Mine Drainage

SRB are commonly employed in AMD treatments where they utilize the sulfate as their terminal electron acceptor reducing it to hydrogen sulfide according to Equation 1.6 while oxidizing the organic matter (Al-Zuhair et al., 2008; Sierra-Alvarez et al., 2006). The produced H2S during sulfate reduction reacts with dissolved metals

Chapter 1 11 1.4.1. Characteristics and diversity of sulfate-reducing bacteria

SRB are a diverse group of bacteria which are abundant in some extreme natural environments and have been isolated from various environments including soil, sediments and non-polluted water (Kakooei et al., 2012; Wargin, 2007). They are considered to be some of the oldest and mostly studied bacterial groups (Wargin, 2007). Their interesting functions in the ecosystem makes them to be widely studied and classified under a number of groups based on their complex characteristics (Castro et al., 2000; Kakooei et al., 2012). They are detected by H2S production in

the environment which has an unpleasant smell and result in black precipitates of ferrous sulfide in the presence of iron minerals (Wargin, 2007). Majority of known cultivable SRB belong to a group of bacteria known as the delta proteobacteria (Baumgartner et al., 2006), which two families of SRB were proposed: the Desulfovibrionaceae and the Desulfobacteriacea. Most abundant SRB isolated from various environments belong to the genera Desulfovibrio and Desulfomicrobium which are mostly the key sulfate reducers in natural anaerobic sediments (Castro et

al., 2000). Castro and co-workers classified SRB into four groups: Gram-negative

mesophilic SRB, Gram-positive spore-forming SRB, bacterial thermophilic SRB and archael thermophilic SRB shown on Table 1.1 (Castro et al., 2000). The SRB classification and representative specie selection was mainly based on their rRNA sequence analysis which provides information about the evolution of the SRB and their distantly related species. The groups were further characterized according to their morphology, nature of motility, GC contend of their DNA%, presence of the Desulfovibrin pigment and cytochromes, complete or incomplete oxidation of acetate and their optimum growth temperatures (Castro et al., 2000). The characterization of SRB based on growth temperature was limited between 25 and 70ºC (Castro et al., 2000; Karr et al., 2005) but some were isolated in temperatures as high as 130ºC in hydrothermal vents (Christophersen et al., 2011; Kallmeyer and Boetius, 2004).

Chapter 1 12 Table 1.1: Classification of representative sulfate-reducing bacteria (Taken from Castro

et al., 2000). Shape Motility GC content of DNA (%) Desulfo-vibrin Cytochr-omes Oxidation of acetate Growth temp (⁰C) Gram-negative mesophilic SRB

Desulfobulbus lemon to rod -/+ 59-60 - b, c, c3 Ia

25-40

Desulfomicrobium ovoid to rod +/- 52-67 - b, c I 25-40

Desulfomonas rod - 66 + c I 30-40

Desulfovibrio

spiral to

vibrioid + 49-66 +/- c3_{, b, c I 25-40}

Desulfobacter oval to rod +/- 44-46 - Cb 20-33

Desulfobacterium oval to rod +/- 41-52 - b, c C 20-35

Desulfococcus

spherical or

lemon -/+ 46-57 +/- b, c C 28-35

Desulfomonile rod - 49 + c3 _{C 37}

Desulfonema filaments Gliding 35-42 +/- b, c C 28-32

Desulfosarcina oval rods to coccoid packages +/- 51 - b, c C 33 Gram-positive spore-forming SRB Desulfotomaculum straight to

curved + 48-52 - b, c I/C most 25-40, rods some 40-65

Bacterial thermophilic

SRB

Thermodesulfobacterium vibrioid to rod -/+ 30-38 - c3, c I 65-70 Archaeal thermophilic SRB Archaeoglobus coccoid +/- 30-38 - n.r.c _{I 64-92} a_{I, incomplete.} b C, complete. c_{n.r., not reported.}

SRB are known to regulate the sulfur cycle in the natural environments and are involved in the organic matter turnover in anaerobic soils and sediments (Castro et

al., 2000; Sánchez-Adrea et al., 2014). Sulfur is one of the most abundant elements

on earth concentrated in the earth’s crust associated with metals occurring as metal sulfides. Sulfur can occur in nine different oxidation states as shown in Figure 1.2 and every oxidation state of sulfur can be acted on by microorganism to change its state of oxidation (Sánchez-Adrea et al., 2014). The final oxidation state of sulfur is sulfate (SO42-) which is the highly reactive with most elements forming compounds

Chapter 1 13

role player in the acidity and toxicity of mine drainages since SO42- concentrations

are elevated (Al-Zuhair et al., 2008; Sánchez-Adrea et al., 2014).

Oxidation and reduction of sulfur and its compounds can be chemically or biologically mediated. Figure 1.2 shows three transformation processes (oxidation, reduction and disproportionation) undertaken by microorganisms in sulfur cycling of the environment. Dissimilatory sulfate reduction which is mediated by SRB in anaerobic conditions where sulfate is reduced to sulfide according to Equation 1.6 (Luptakova et al., 2012; Sánchez-Adrea et al., 2014). Dissimilatory sulfate reduction is mediated by some SRB where they convert the elemental sulfur into hydrogen sulfide. Assimilatory sulfate reduction which all SRB are capable of is one of the sulfate reduction strategies since the reduced sulfide is assimilated in the bacterial biomass, proteins and amino acids (Christophersen et al., 2011). Sulfide oxidation of reduced sulfur species back to sulfate by sulfide-oxidizing bacteria (SOB) such as methanogenic, lithotrohic and phototrophic bacteria which uses mostly ferric iron and oxygen as their electron acceptors. To continue the cycle from sulfate to sulfide, sulfate reduction by SRB is required to transform sulfur species (Dahl et al., 2005; Foti et al., 2007). Lastly, disproportionation of sulfur species where coupled oxidation and reduction of sulfur compounds such as thiosulfide (S2O32-), sulfite (SO32-) and

elemental sulfur (S80) which are mostly present as intermediate compounds during

sulfate reduction to hydrogen sulfide (Sánchez-Adrea et al., 2014).

Figure 1.2: Biological sulfur transformations in natural and artificial environments. (Taken from Sánchez-Adrea et al., 2014).

Chapter 1 14 1.4.2. Factors affecting biological sulfate reduction in AMD

Function of SRB in AMD remediation is simply to convert sulfate into sulfide reducing the SO42- concentrations in the drainage simultaneously utilizing protons hence

neutralizing the drainage. The biogenic sulfide (S2-) reacts with dissolved metals such as Fe2+, Pb2+, Cu+, Cd2+, Ni4+ and Zn2+ forming insoluble precipitates which is an indirect metals removal from the drainage by SRB (Sánchez-Adrea et al., 2014). This reversal process of AMD formation is largely dependent on the environmental conditions which SRB operates in. The first priority for a successful AMD remediation by SRB is an anoxic to anaerobic condition which will allow maximum sulfate reduction to take place with less to no sulfide oxidation. Then the process should be optimized with a suitable carbon source and electron donor selection. Other non-compulsory requirements for successful sulfate reduction are optimum pH and temperature which differs according to various SRB groups (Castro et al., 2000; Kakooei et al., 2012). Concentrations of sulfate, sulfide and different metals can also negatively affect the sulfate-reduction capacity of SRB (Russel et al., 2003).

1.4.2.1. Electron donor and carbon source selected for SRB growth

Due to low concentrations of organic matter especially carbon based minerals in the AMD, additional carbon source and electron donor are required when growing SRB. SRB obtain their energy through the oxidation of organic compounds according to Equation 1.6 (Rzeczycka and Blaszczyk, 2005). One factor that limits the microbial sulfate reduction and metals removal from AMD is the supply of energy and carbon source used by the SRB (Russel et al., 2003). Various carbon sources had been used in the isolation and growth of SRB and sodium lactate was reported to be more preferred donor by SRB (Wargin, 2007). Variety of electron donors such as alcohols, aromatic compounds, fatty acids and hydrogen are usable by SRB to reduce sulfate but utilization of these various donor sources by microorganisms does not yield similar energy levels seen by growth rates and biomass production of the SRB (Karr

et al., 2005). Utilization of organic acids (lactic, pyruvic, formic and acetic) and

alcohols (ethanol and propanol) by SRB as electron donors have been studied widely and used as a classification scheme dividing SRB based on type of organic

Chapter 1 15

substrate used and if the substrate is completely or incompletely oxidized and degraded (Castro et al., 2000; Rzeczycka and Blaszczyk, 2005). The most important factors to consider when choosing the suitable carbon source and electron donor for a specific group of SRB are the ability of SRB to utilize the substrate, the suitability of the substrate for the treatment system. This is coupled to the amount of sulfate to be reduced through the use of the selected substrate (Kaksonen and Puhukka, 2007). Wide range of electron donor options for SRB reveals the less fastidious requirements of nutrients to grow SRB but definitely a suitable one is required to obtain desired sulfate reduction results using SRB (Rzeczycka and Blaszczyk, 2005; Zagury et al., 2006).

1.4.2.2. The effect of temperature on SRB

SRB can be classified into three groups based on the temperature at which they optimally reduce sulfate. First being the mesophiles which grow at temperatures less than 40ºC such as Desulfovibrio sp. which is known to optimally grow at temperatures between 25 and 40ºC (Sawicka et al., 2012). Secondly are the moderate thermophiles which are SRB growing optimally at temperatures between 40 and 60ºC such as Desulfotomaculum sp. (Castro et al., 2000; Fichtel et al., 2012). The last group are the thermophilic SRB which can grow at temperatures from 60 till 110ºC such as Archaeoglobus sp. which was reported to have been isolated from hydrothermal vent chimney and can grow at 88ºC (Belkin et al., 1985). Although these groups have been studied, the detection and successful isolation of SRB from extreme conditions like frozen lakes in the Arctic where the temperature ranges between -1 to 8ºC has been reported and the isolated SRB are called psychrotolerant SRB. The highest sulfate reduction percentages obtained by the psychrotolerant/mesophilic SRB was reported to be between 20 to 50% at temperature ranges of 0 to 40ºC (Moosa et al., 2005; Sawicka et al., 2012). SRB successfully used to reduce sulfate at 102ºC was isolated from a thermal vent with the highest recorded temperature of 130ºC. These bacteria showed typical characteristics of archaea growing in extremely hot conditions (Fichtel et al., 2012; Kallmeyer and Boetius, 2004). Even though SRB have been isolated from various habitats with extreme temperatures, that does not mean they can optimally reduce

Chapter 1 16

sulfate completely to sulfide in these environmental conditions similar to those of their isolation habitats hence application of SRB in bioremediation require optimization. Biological sulfate reduction is possible and efficient at temperatures above 30⁰C but the SRB activity slows down at low temperatures below 10⁰C (Kakooei et al., 2012).

1.4.2.3. The effect of pH on sulfate reduction

Most of the known SRB have been isolated from AMD with pH values lower than 4 indicating their tolerance to acidic conditions and most likely high metals concentrations (Hard et al., 1997). The presence of SRB in acidic environments does not mean their condition of origin is their optimum for sulfate reduction. Most of the studied SRB have been reported to grow optimally at pH between 6 and 8 which is the desired pH level when treating acidic drainages (Sánchez-Adrea et al., 2014). SRB that are susceptible to low pH tends to result in minimal sulfate reduction and low biomass yields in these unfavourable conditions. Acidic environments affect SRB negatively compared to neutral and slightly alkaline environments because most of organic acids do not dissociate in low pH environments hence limiting carbon supply to the SRB for growth (Hard et al., 1997; Kaksonen and Puhakka, 2007). Resistance of acidophilic SRB to low pH conditions is due to the fatty acids contained in their cell walls which are not vulnerable to the acidic medium and hence they thrive in these harsh conditions. This is in relation to the neutrophilic SRB with mostly acid permeable cell walls that allow acids to diffuse through the membrane resulting in the acidification of the cytoplasm which could be lethal to microbial cells. This could all lead to low sulfate reduction in the acidic environments. To avoid such toxicity, consortia containing both acidophilic and neutrophilic SRB can be used or the AMD being treated could be pre-treated with neutralizing agents to raise the pH prior to introduction to the SRB inoculum (Fortin et al., 1996; Postgate, 1984). The other vital purpose of keeping pH above 5 is to keep the hydrogen sulfide (HS-) in the solution form for the reaction with dissolved metals that enhances metals sulfide precipitation which is slowed down at low pH because of the sulfide being in a gas form (H2S) hence unavailable to dissolved metals (Fortin et al., 1996; Kaksonen and

Chapter 1 17 1.4.2.4. Sulfate and sulfide concentrations affecting sulfate

reduction

Presence and activity of SRB in the environment is detected by sulfide production which is noted by a pungent odour similar to that of rotten eggs. Biogenic hydrogen sulfide which forms according to Equation 1.7 can be present in three different forms (H2S, HS- and S2-) in either liquid or gaseous phase (Kaksonen and Puhukka, 2007).

The concentration of each form depends on the reversible chemical equilibria represented by Equation 1.8 and 1.9 at temperatures between 25ºC and 30ºC (Kaksonen and Puhukka, 2007).

Equation 1.8: 𝐻2𝑆 ↔ 𝐻𝑆− + 𝐻+

Equation 1.9: 𝐻𝑆− ↔ 𝑆2− + 𝐻+

The form at which sulfide occur depends on the pH of its environment. At pH values below 6, hydrogen sulfide is mostly in an undissociated gaseous form which is toxic to most SRB at high concentrations. The undissociated H2S affect the metabolic

coenzymes and denature proteins by precipitating metal ions in the active sites within the bacterial cells. In terms of metals precipitation by the biogenic sulfide, the gaseous form easily diffuses out of the solution without much interaction with dissolved metals in cases where opened systems like wetlands are employed. In closed anaerobic systems, the gaseous hydrogen sulfide creates pressure which leads to partial dissociation and dissolves in the medium which will later negatively affect the bacterial cells (Kaksonen and Puhukka, 2007; Postgate, 1984).

At pH values between 6 and 8, the H2S dissociates according to Equation 1.8 to

yield hydrogen sulfide (HS-) which is mostly in a liquid form available to dissolved metals. The dark grey shaded portion of the graph in Figure 1.3 highlights the pH range where the dissociation of H2S to HS- takes place and shows dominance

beyond that point. The higher pH induces metals precipitation as they react with sulfides present in the medium. The sulfide in a form HS- is not directly toxic to SRB but indirectly affect their metabolisms by precipitating metals like iron which are needed by some groups of SRB. The biologically induced metal sulfide precipitates

Chapter 1 18

are a positive sign of successful sulfate reduction and metals removal from the medium which are part of the remediation strategies by SRB.

Figure 1.3: Sulfide speciation as a function of pH at 25ºC. (Taken from Kaksonen and Puhukka, 2007).

1.4.2.5. Effect of high metals concentrations on sulfate-reduction

Metals such as iron, zinc, copper and lead are toxic to SRB at high concentrations. Metal sulfide precipitates that form during sulfate reduction are not directly toxic to SRB but can block access of substrates and nutrients to SRB as they accumulate around cells in the biofilm. Some of the metals contained in AMD can pass through the cell wall and membranes of the cells and precipitate with the intra-generated H2S

resulting in precipitates within the cell which become toxic as they accumulate within the cells (Utgikar et al., 2002).

Metals can affect SRB by exerting their toxicity on the biological systems of the cells by altering multiple biochemical pathways simultaneously. According to Harrison and co-workers (2007), metals can affect SRB biochemically by substitutive metal-ligand binding where one metal ion replaces another at the binding site of a specific biomolecule which destroys the biological function of the target molecule. The second metal effect mechanisms is the covalent or ionic redox (reduction-oxidation) reactions of metal ions with cellular thiols (R-SH) such oxyanions which liberate

Chapter 1 19

reactive oxygen species like superoxide (O2.-) as by-products which are toxic to

growing bacteria. Some metals affect the membrane transport process where the toxic metals occupy the binding sites inhibiting biochemical reactions in the cell to proceed normally (Harrison et al., 2007). Most of the bacteria isolated from streams and sediments with toxic metals have thrived through the conditions by making biofilms which protect most of the inter biofilm bacterial cell which tends to develop resistance against the toxic metals (Harrison et al., 2007; Teitzel and Parsek, 2003).

Chapter 1 20

1.5. Conclusions

Acid mine drainage is generated by oxidation of metal sulfide minerals like pyrite, galena, sphalerite and chalcopyrite contained in the earth’s crust referred to as minerals host rock. Oxidation of these minerals takes place during anthropogenic activities, principally mining. The oxidation process of these minerals results in low pH conditions with high concentrations of dissolved sulfate and metals. Microorganisms such as Acidithiobacillus ferrooxidans catalyses the AMD generation rates up to 104 to 106 times. Depending of the chemical contents of the drainage, high pH conditions of some drainages are possible with low metals concentrations but high sulfate concentrations and this drainage is called non-acid mine drainage (NMD) or alkaline mine drainage.

Biological and chemical treatment options are available to aid in AMD remediation but most of the systems are expensive to implement and maintain. AMD treatment, especially aided by biological sulfate reduction can be enhanced if SRB are characterized. A cost effective, efficient and reliable system is required for both active and abandoned mines generating million litres of AMD contaminated water. Due to the high cost demands of neutralizing chemicals like calcium carbonate, sodium carbonate and bicarbonate to treat AMD, various passive biological systems have also been developed to treat AMD with lower costs. The reversal process of AMD formation can be mediated by application of microorganisms such as sulfate reducing bacteria (SRB).

SRB are anaerobic bacteria that utilizes sulfate as their terminal electron acceptor to oxidize the organic matter while releasing hydrogen sulfide. The biogenic hydrogen sulfide will react with dissolved metals and precipitate them which will lead into the elevation of pH as sulfate concentrations gets reduced. The use of SRB in AMD or NMD interaction requires background knowledge of the SRB and factors that affect sulfate reduction like carbon source, temperature, pH changes and metals concentration require constant monitoring to obtain best remediation results. The SRB mediated metal precipitation as part of mitigating AMD or NMD also require attention to evaluate the limitations metal concentrations exert on the bacteria. The identity and characterization of phase minerals present in biogenic precipitates still require further exploration which is the partial focus in this study.

Chapter 1 21

1.6. References

 Akcil, A. and Koldas, S. (2006) Acid mine drainage (AMD): causes,

treatment and case studies. Journal of Cleaner Production 14: 1139 – 1145.

 Al-Zuhair, S., El-Naas, M. and Al-Hassani, H. (2008) Sulfate inhibition effect

on sulfate reducing bacteria. Journal of Biochemical Technology 1(2): 39 – 44.

 Atkins, A. S. and Pooley, F. D. (1982) The effect of Bio-mechanism on acidic

mine drainage in coal mining. International Journal of Mine Water 1: 31 – 44.

 Baker, B. J. and Banfield, J. F. (2003) Microbial communities in acid mine

drainage. Federation of European Microbiological Societies 44: 139 – 152.

 Baker, B.J., Moser, D.P., MacGregor, B.J., Fishbain, S., Wagner, M., Fry, N.K., Jackson, B., Speolstra, N., Loos, S., takai, K., Lollar, B.S., Fredrickson, J., Balkwill, D., Onstoff, T.C., Wimpee, C.F. and Stahl, D.A.

(2003) Related assemblages of sulphate-reducing bacteria associated with ultra-deep gold mines of South Africa and deep basalt aquifers of Washington State. Environmental Microbiology. 5: 267-277.

 Baumgartner, L. K., Reid, R. P., Dupraz, C., Decgo, A. W., Buckley, D. H., Spear, J. R., Przekop, K. M. and Visscher, P. T. (2006) Sulfate reducing

bacteria in microbial mats: Changing paradigms, new discoveries.

Sedimentary geology 185: 131 – 145.

 Belkin, S., Wirsen, C. O. and Jannasch, H. W. (1985) Biological and

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